7. AQUATIC LIFE

7.1 Effects

7.1.1 Freshwater environment

(a) Lethal and acute toxicity to aquatic animals

The lethal toxicity (2- to 30-d LC₅₀) of PCBs to freshwater organisms varies with several factors which include the PCB formulation, the organism species and stage of development, and the test conditions employed (e.g., length of exposure, static versus flow-through tests, etc.) (Table 10; Figures 2 through 6). Aroclors containing 42 to 54% chlorine appear to be the most toxic formulations of PCBs; for instance, a static test with Daphnia magna exposed to Aroclors A-1221, A-1242, A-1248, A-1254, A-1260, A-1262, and A-1268 displayed 504-h LC50s of 180, 67, 25, 31, 36, 43, and 253 µg/L, respectively (Nebeker and Puglisi, 1974; Table 10). Similar observations can be made from tests with rainbow trout (Oncorhynchus mykiss formerly classified as Salmo gairdneri ) exposed to several PCB formulations in a continuous flow-through system (Johnson and Finley, 1980; Mayer et al., 1977; Table 10).

The toxicity of PCBs is more severe when aquatic organisms are exposed to them in a continuous flow-through system rather than in a static system; also, longer exposure periods yield lower LC₅₀s (Nebeker and Puglisi, 1974; Stalling and Mayer, 1972; Table 10). The lowest concentrations causing lethal toxic effects were: 23 µg/L (720-h LC₅₀, Aroclor 1016), 5 µg/L (240-h LC₅₀, Aroclor 1242), 2.6 µg/L (336-h LC₅₀, Aroclor 1248), 0.45 µg/L (504-h LC₅₀, Aroclor 1254), and 3.3 µg/L (720-h LC₅₀, Aroclor 1260). In general, the data in Table 10 and Figure 2 through 6 appear to suggest that invertebrates and fish are equally sensitive to PCBs.

(b) Sublethal and chronic toxicity to aquatic organisms

The information obtained from the literature on sublethal and chronic effects of PCBs to freshwater aquatic organisms is tabulated in Tables 11 and 12. The data were also plotted in Figures 7 and 8. The lowest observed adverse effect level for algae was 1.0 µg/L Aroclor 1242, but, in general, aquatic animals appear to be more sensitive to PCBs than algae and plants. The minimum concentration of commercial PCBs (e.g., Aroclors A-1242 and A-1254) causing chronic effects (e.g., inhibitory effects on the ATPase activity of brain and kidney tissues of fathead minnow) was recorded to be 0.31 µg/L by Koch et al., 1972 and Cutkomp et al., 1972. These investigators also found that the effect on ATPase activity of the fish tissues was not consistent with the toxicant concentration; for instance, in several instances the enzyme activity at a higher exposure level of 8.3 µg/L of the toxicant was comparable to the control. In a 30-d chronic exposure to 0.1 µg/L Aroclor 1248, DeFoe et al. (1978) found that both the first- (F1) and the second-generation (F2) fathead minnow (P. promelas ) fry and larvae survival, weight, or length were not affected.

* S = Static, CF = Continuous flow, M = Measured concentration; U = Unknown

Figure 2

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Figure 7

Figure 8

Addison et al. (1978) found that trout (S. fontinalis ) fed Aroclor 1254 to produce a tissue (fillet) concentration of 39 µg PCB/g wet weight showed a significant increase in the activity of ethoxycoumarin O-deethylase (ECOD). During the feeding experiment, one fish out of the six died after 17 days on the PCB diet. Also, there was substantial variation in the sensitivity of various MO (mono-oxygenase) activities to xenobiotics. For instance, (i) EROD activity increased in rainbow trout, carp, and channel catfish treated with single intra-peritoneal injections of 1.0 µg Aroclor 1254/g body weight (bw) (Melancon and Lech, 1983; Ankley et al., 1986); (ii) ECOD activity in rainbow trout was also increased by a dose of 1.0 µg Aroclor 1254/g body weight. However, ECOD activity in catfish did not increase until treated with a dose of 10 µg Aroclor 1254/g bw, and, in carp, it was not stimulated by doses as high as 200 µg Aroclor 1254/g bw (Melancon et al., 1981; Melancon and Lech, 1983; Ankley et al., 1986).

Using groups of eight rainbow trout (O. mykiss ) fed rations containing 0, 1, 10 and 100 µg/g Aroclor 1254 over a period of 330 days, Nestel and Budd (1974) found pathological changes in the kidneys of 13 fish. The greatest number of cases occurred in fish at the 10 µg/g concentration. Fish fed 1 and 10, and 100 µg/g diet had mean residues of 1.4, 2.3, and 80.1 µg PCB/g (wet weight), respectively. These investigators also noted that the rainbow trout survived an oral intake of PCBs at the rates of 0.04 µg, 0.4 µg, and 4.0 µg per g body weight (bw) for 330 days. The results of the Nestle and Budd (1974) study combined with those of Melancon and Lech (1983) above, appear to suggest that the induction of EROD activity at 1.0 µg PCB/g bw may not necessarily mean initiation of toxic effects (including death).

(c) Toxicity of PCB isomers or congeners

Industrial PCBs are mixtures of several isomers and congeners. The structure-activity relationships have shown that several PCB congeners are similar to 2,3,7,8-TCDD in structure and in induction of aryl hydrocarbon hydroxylase (AHH) and ethoxyresorufin O-deethylase (EROD) activity in rat hepatoma cells. Based on these structure-activity relationships, it was found that the true coplanar PCBs (#77, #126, and #169) and coplanar PCBs which are additionally halogenated in meta or para positions, are among the most toxic PCBs. Among 15 congeners tested, 3,3',4,4',5 pentachlorobiphenyl (PCB #126) was the most potent inducer of the AHH and EROD activities and was followed in the induction response by 3,3',4,4',5,5' hexachlorobiphenyl (PCB #169) and 3,3',4,4' tetrachlorobiphenyl (PCB #77) ( Table 6; Safe, 1987; Sawyer and Safe, 1982).

The congener-specific toxicity data for aquatic life are limited, especially with respect to the relatively more toxic coplanar congeners. Also, the available data show that a non-coplanar PCB congener (e.g., PCB #52) can be more severely toxic to amphipod Hyalella. azteca than a coplanar congener (e.g., PCB #77) (Table 13; Figure 9). The most severe chronic toxicity to a PCB congener was displayed by D. pulicaria exposed to PCB #4 at 0.05 µg/L (Bridgham, 1988; Figure 9).

7.1.2 Marine environment

(a) Lethal and sublethal toxicity

The lethal and sublethal toxic effects of PCB formulations on marine organisms are shown in Tables 14 and 15 and Figures 10 through 13. The data suggest that PCBs are as toxic to marine organisms as they are to freshwater organisms. In flow-through chronic tests with Aroclor 1254, Hansen et al. (1973, 1974) found that the survival of fry and the hatching of embryos from exposed sheepshead minnow adults were affected by 0.1 µg/L PCB (measured concentration). Adverse effects were also observed on the growth of a marine diatom (Rhizosolenia setigera ) exposed to 0.1 µg Aroclor 1254/L; the growth was more severely reduced at a lower temperature (10 C) than at a higher temperature (15 C) (Fisher and Wurster, 1973) .

Cosper et al. (1987) found that the marine diatom Ditylum brightwellii, pre-treated with sublethal concentrations of 10 to 30 µg/L over a period of 30 days, developed a resistance to PCBs. The PCB-resistant strain exhibited greater tolerance to PCB than the PCB-sensitive strain under all environmental conditions. However, PCB resistance decreased the tolerance of the strain to lower salinities and nitrogen limitation, but increased its tolerance to lower temperatures.

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Figure 13

Gruger et al. (1977) reported induction of hepatic aryl hydrocarbon hydroxylase (AHH) activity in coho salmon (Oncorhynchus kisutch ) exposed to 1.0 µg/g of Aroclor 1242 in diet (wet weight). No adverse effects of the concentration of PCBs in fish, however, were reported by the investigators. These observations concur with those reported for freshwater fish in section 7.1.1 (b).

(b) Toxicity of PCB isomers or congeners

No congener-specific toxic reactions to PCBs were found in the literature for marine organisms.

7.1.3 Bioaccumulation of PCBs in aquatic organisms

Accumulation of PCBs in aquatic organisms results from their uptake from food and water. This type of residue buildup in aquatic organisms through water and food is appropriately defined in terms of bioaccumulation factor (BAF). On the other hand, accumulation of contaminants in tissues of aquatic organisms from water alone is defined in terms of bio-concentration or bio-concentration factor (BCF). Table 16 contains the results of several freshwater and marine residue studies.

Fish exposed to PCBs in water alone have shown an excellent correlation between BCF (i.e, contaminant concentration in fish ÷ contaminant concentration in water) and the contaminant's octanol-water partition coefficient (Kow) (Mackay, 1982; Oliver and Niimi, 1983, 1985). One such model (Mackay, 1982) was used by Oliver and Niimi (1988) to study accumulation of PCBs in salmonids in the Lake Ontario ecosystem. These investigators concluded that: (a) contaminated food was a major source of the PCB residues in fish, (b) PCB uptake from water alone would underestimate fish residues by at least a factor of 5 (BCF of 780 000 was estimated for salmonids in the Lake Ontario ecosystem), and (c) a bioaccumulation factor of 3 900 000 (the highest value for freshwater aquatic life in Table 16) was found for total PCB in fish.

In most cases, the results shown in Table 16 (except Oliver and Niimi, 1988) were obtained from organisms exposed to the contaminants in water alone under controlled conditions; hence, they represent bio-concentration factors (BCFs) rather than bio-accumulation factors (or BAFs). The maximum BCF for marine life was found to be 13 000 000 (expressed on lipid basis) for Northern anchovy larva, which would translate to 975 000 when expressed on whole fish basis (lipid content of the fish was 7.5%) (Scura and Theilacker, 1977).

It is evident from Table 16 and the above discussion that PCBs in freshwater and marine organisms bio-concentrate to a similar degree.

7.1.4 Sediment toxicity

Sediments act as a sink for contaminants such as PCBs. The bio-availability of sediment-associated contaminants is central to whether the toxic compounds (e.g., PCBs) present in the sediment will have deleterious effects on aquatic species or will become part of the food chain.

In studying PCB availability to marine animals from the spiked sediments (containing 1.0 µg/g PCBs), McLeese et al. (1980) found that the bioaccumulation of PCBs in polychaetes (Nereis virens) and shrimp (Crangon septemspinosa ) was directly related to the concentration of the contaminants in the sediment and inversely related to animal size; the sediment produced no toxicity among the test organisms. A bioaccumulation factor ranging from 1.6 to <0.2 was found for polychaetes (Nereis virens) exposed to natural sediment containing PCBs; accumulation in clams (Mercenaria mercenaria) and shrimp (Palaemonetes pugio), on the other hand, was much less (Rubinstein et al., 1983). Stein et al. (1987) compared the accumulation of PCBs in a benthic fish (English sole, Parophrys vetulus) exposed for up to 108 days to a test (2.2 µg PCBs/g dry weight) and a reference (PCBs at non-detectable level of <4.9 ng/g dry weight) sediment. English sole exposed to the test sediment had a hepatic concentration of 1.4 ± 0.6 µg PCBs/g wet weight, which was eight times greater than that for the reference sediment.

The bio-availability of PCBs from sediment is a function of several factors including concentration of the contaminants in the sediment, exposure time, species, and sediment characteristics (e.g., particle size, organic carbon, etc.). Lynch and Johnson (1982) demonstrated that increased organic matter content and a larger particle size of the natural sediment reduced the concentration of 2,2',4,4',5,5'-hexachlorobiphenyl (hexaCBP; PCB #153) in water and in the freshwater benthic amphipod (Gammarus pseudolimnaeus). Their results also indicated that substrate characteristics, especially the organic matter content, control the availability of hexaCBP to the overlying water and, in turn, to the aquatic organisms.

Tatem (1986) exposed freshwater prawns (Macrobrachium rosenbergii ) and clams (Corbicula fluminea ) to sand and sediment mixtures, containing 10%, 50%, and 100% of the sediment 5-80 (61.1 µg PCBs/g) in one set and the same proportions of the sediment 11-80 (2.3 µg PCBs/g) in the other. (The sediments 5-80 and 11-80 were dredged materials obtained, respectively, in May and November 1980 from one site in Sheboygan River, Mississippi). The mixtures containing 10% sediment produced the maximum bioaccumulation of PCBs in the organisms. The maximum values for bioaccumulation factor (BAF) were determined to be 0.9 (for Aroclor 1242 in sediment 5-80) and 2.4 (for Aroclor 1254 in sediment 5-80) for prawns, and 12.52 (for Aroclor 1242 + Aroclor 1254 in sediment 11-80) for clams.

To assess the likelihood or potential for adverse biological effects of sediment-associated toxicants (e.g., PCBs) to biota, Long and Morgan (1990) assembled data derived from a wide variety of methods and approaches (e.g., the equilibrium partitioning approach, the spiked-sediment bioassay approach, etc.). The data were evaluated to identify informal guidelines for use for sediments under the National Status and Trends Program. The contaminant concentrations observed or predicted by various methods to be associated with biological effects were sorted. The lower 10th percentile (i.e., Effects Range-Low or ER-L, defined as a concentration at the low end of the range in which effects had been observed) and median (i.e., Effects Range-Median or ER-M, defined as a concentration approximately midway in the range of reported values associated with biological effects) values were identified. These investigators obtained ER-L and ER-M values, respectively, of 0.05 µg PCBs/g sediment and 0.4 µg PCBs/g sediment from the data used. The ER-L obtained by Long and Morgan was supported by the apparent effects threshold (i.e., AET, defined as the sediment concentration of a contaminant above which statistically significant biological effects, such as depression in abundance of benthic infauna or elevated incidence of mortality in sediment toxicity tests, are always expected) for bivalve larvae for San Francisco Bay. Furthermore, the ER-M value was observed to be similar to the mean level (0.368 µg PCBs/g) in the Commencement Bay (Washington) sediment samples which were highly toxic to oyster larvae and the mean concentration (0.4 µg PCBs/g) in southern California sediments with moderate species richness.

7.2 Criteria from the literature

7.2.1 Ambient water

Criteria, objectives, and standards to protect aquatic life from the harmful effects of PCBs are shown in Table 17. The recommended criterion varied with jurisdiction. The most stringent criteria (0.0079 - 0.79 ng/L) were proposed by the U.S. Environmental Protection Agency (1980) for the protection of humans against cancer risk from consuming PCB contaminated water and aquatic organisms.

In developing its criterion based on bio-magnification of PCBs in fish, the Ontario Ministry of the Environment assumed a tolerance level of 2 µg/g wet weight in the edible tissue. This upper limit of 2 µg PCBs/g wet weight was established by Health and Welfare Canada as an action level for the sale and export of fish for human consumption. A similar approach was used by the International Joint Commission (1977), but it recommended a concentration of total PCBs in fish tissue (whole fish) not exceeding 0.1 µg/g wet weight to protect fish-consuming birds and animals. The recommended concentration of 1 ng PCBs/L in water by IJC was based on (i) Platonow and Karstad (1973) studies on commercial ranch mink where the lowest dietary concentration observed to cause deleterious effect was 0.64 µg/g Aroclor 1254, (ii) an application of a safety factor of 5, and (iii) a bio-concentration factor of 100 000. The criterion of 1 ng/L was also recommended by CCREM, Ontario, Indiana, Ohio, and Pennsylvania.

The U.S. EPA (1980) criteria for ambient waters were developed to protect freshwater aquatic life, marine aquatic life, and human health. The concentrations of 14 ng/L for freshwater and 30 ng/L for marine environments were considered too high by the EPA as they were based on bio-concentration factors measured in laboratory studies (BCFs for fish from field studies are at least 10 times higher). It was, therefore, recognised that these criteria would provide adequate protection only against acute effects of PCBs. The Province of Manitoba adopted the U.S. EPA criterion of 14 ng/L for freshwater.

For the protection of human health from potential carcinogenic effects of PCBs ingested from the use of contaminated water and contaminated aquatic organisms, the ambient water concentration of PCBs should be zero according to the U.S. EPA (1980). The U.S. EPA, however, recognised that this level may not be attainable at this time. As a result, PCB levels in water were recommended considering the increased risk of developing cancer in a lifetime. Based on the consumption of 2 L/d of contaminated water and 6.5 g/d of fish taken from the contaminated water, it was recommended that PCB levels in water should not exceed 0.79 ng/L, 0.079 ng/L, and 0.0079 ng/L for increased cancer risks of 1 in 100 000, 1 in 1 000 000, and 1 in 10 000 000, respectively.

Recently, the Canadian Council of Environment Ministers (CCME), formerly known as CCREM, recommended a concentration of 10 ng/L in saltwater to protect marine aquatic life (CWQG, 1991).

7.2.2 Fish and/or shellfish

The International Joint Commission (1977) recommended that the concentration of total PCBs in fish tissue (whole fish) should not exceed 0.1 µg/g wet weight to protect fish-consuming birds and animals. Health and Welfare Canada (1975) recommended the maximum tolerance level of 2 µg/g wet weight in fish (edible portion) to protect humans. This guideline was based on several factors which include maximum residue levels in all foods other than fish, economic impact to the fishing industry, and the recommended 'tolerable daily intake' of 1.0 µg/kg body weight/d for PCBs in Canada (Grant, 1983). Similar levels ( 2µg PCBs/g wet weight) in fish were adopted by the Ontario Ministry of Environment (1985) and the U.S. Food and Drug Administration (1984).

The objective of 0.5 µg/g (wet weight) PCBs in fish tissue was recommended for the Fraser River and Burrard Inlet in British Columbia (Swain and Holms, 1984; Nijman and Swain, 1990).

7.2.3 Sediments

The sediment criteria from various jurisdictions are shown in Table 18. A comparison among the jurisdictions is difficult to make since the guidelines are not expressed in consistent units (i.e., they are not normalised to organic carbon content). The lowest value

* Dry weight basis

of 0.03 µg/g for total PCBs is an objective for the lower Fraser River and Estuary, Boundary Bay, and Burrard Inlet, set by the British Columbia Ministry of Environment based on measurements for uncontaminated sites. The highest value of 10 µg/g total PCBs is set by Region V of the U.S. EPA (1977); this guideline will likely be superseded by the U.S. EPA (1989) sediment criteria currently under development.

The State of Washington criteria are given in terms of Screening Level Concentration (SLC, i.e., estimated highest concentration of a non-polar contaminant that co-occurs with approximately 95% of the infauna) and AET (Table 18). The recommended SLC (0.1 µg PCBs/g sediment) and AET (0.13 µg PCBs/g sediment) for Puget Sound are similar, but twice the Effects Range-Low (ER-L) value of 0.05 µg PCBs/g sediment obtained by Long and Morgan (1990) or the AET recommended for San Francisco Bay (see Section 7.1.4). Since organic carbon content of the sediments was not stated in these references, a direct comparison is difficult to make between the criteria recommended by the State of Washington (Table 18) and the ER-L and ER-M proposed by Long and Morgan. Note that the availability of PCBs in sediment is strongly dependent upon its organic carbon content.

7.3 Recommended Criteria

7.3.1 Freshwater and marine aquatic life

For the protection of freshwater and marine aquatic life and consumers of fish and shellfish (e.g., wildlife), it is recommended that the total PCB concentration in water should not exceed 0.1 ng/L. Additionally, it is recommended that the concentration of some selective PCB congeners (e.g., PCB congeners 77, 105, 126, and 169) should not exceed levels shown in Table 19.

The recommended guidelines by CCREM (1987) and CCME (CWQG, 1991) for freshwater and marine water, respectively, are 1 ng PCBs/L and 10 ng PCBs/L.

7.3.2 Fish and shellfish

To protect wildlife dependent on aquatic life for food, it is recommended that the concentration of PCBs in fish and/or shellfish should not exceed 0.1 µg/g (wet weight).

* based on 2,3,7,8-TCDD criterion in water of 0.1 pg/L and the maximum value of the TEF-range shown in Table 6.
+ From Oliver and Niimi, 1988; Hansen, 1987; Kannan et al., 1988; Bush et al., 1985
++ The absence of data does not necessarily mean the congener is absent; it may be below the reliable detection limit or the standard may not have been available.

To protect human consumers from PCB residue in aquatic life, it is recommended that the concentration of PCBs in the edible portion of fish and/or shellfish should not exceed 2.0 µg/g (wet weight).

7.3.3 Sediments

To protect aquatic life and consumers of aquatic life (e.g., wildlife), it is recommended that the concentration of PCBs in freshwater and marine sediments containing 1% organic carbon should not exceed 0.02 µg/g sediment (dry weight) (or 2 µg/g organic carbon, when expressed on an organic carbon basis).

7.3.4 Application of criteria

In Section 7.3.1, the recommended criteria for PCBs for fresh and marine waters are given in terms of total PCB concentration as well as some selected PCB congeners. The measurement of total PCB concentration will provide protection against the effects that may be caused by most of the congeners listed in Table 19, as long as the criterion for total PCB is met. However, the criteria recommended for congeners #77, #105, #126, and #169 are more stringent than the total PCB criterion of 0.1 ng/L. Since these coplanar congeners (i.e., PCB #77, #105, #126, and # 169) are present in most of the commercially available PCB formulations, it is recommended that they should also be measured to ensure that the PCB criteria in water are met in all respects. Both the total and congener-specific PCB criteria should be met.

7.4 Rationale

7.4.1 Freshwater and marine aquatic life

The criteria to protect freshwater and marine aquatic life from accumulating undesirable levels of PCBs in their tissue are the same, and were based upon the information presented in Platonow and Karstad (1973), Section 7.1.3, and Table 16. Considering a maximum acceptable toxicant concentration 0.1 µg PCBs/ g (wet weight) in fish consumed by wildlife (see Section 8.4), and a bio-concentration factor of ~1 000 000 for fish from water alone (which appears to be the same for both freshwater and marine environments - see Section 7.1.3), it was calculated that the concentration equal to or lower than 0.1 ng PCBs/L in water (i.e., 0.1 µg/g ÷ 1 000 000 = 0.1 ng PCBs/L) should protect fish from excessive accumulation PCBs in their tissues.

The criteria for toxic PCB congeners were based on toxic equivalent factors shown in Table 6. It was assumed that the maximum level for 2,3,7,8-TCDD in water is not to exceed 0.1 pg/L, the recently set water quality criterion by the Ontario Ministry of Environment (Lupp and McCarty, 1989). The congener-specific criteria (based on toxic equivalent factors) and concentrations in water from various sources are shown in Table 19.

The criterion of 0.1 ng/L total PCBs recommended in this document either exceeds slightly or is several orders of magnitude above the recommended concentrations for congeners #77, #105, #126, and #169 in Table 19. Although these congeners are very toxic, their concentrations are generally low in natural waters as well as in the most common PCB formulations (Table 19). For instance, PCB # 105 constituted, at the maximum level, 1.3% of the total PCB in water. At the recommended level of 0.1 ng/L total PCB, the concentration of congener #105 would be 0.006 ng/L which is at least an order of magnitude lower than the congener criterion derived in Table 19. However, for congener #77, the concentration in water may exceed the criterion for total PCBs.

No rationale, in clear terms, was provided by CCREM (1987) in setting their criterion of 1 ng PCBs/L for the protection of freshwater aquatic life. However, the interim guideline of 10 ng PCBs/L for the protection and maintenance of marine aquatic life was based on the application of a safety factor of 0.1 to the lowest observed effect level (0.16 µg/L) observed with C. variegatus in a 21-d study (Schimmel et al., 1974). Bioaccumulation models, as in this document, were not used by the CCME (CWQG, 1991)

7.4.2 Fish and shellfish

The criterion (0.1 µg PCBs/g - wet weight - see Section 8.4) for fish and shellfish was based on adverse effects in mink fed 0.64 µg PCBs/g in their diet (containing meat from cows which had been fed Aroclor 1254 (Platonow and Karstad, 1973). The maximum residue level (i.e., 0.1 µg PCBs/g - wet weight) recommended in this document for the protection of wildlife is one-sixth the lowest concentration (in the mink diet) used in the Platonow and Karstad study.

The International Joint Commission (1977) guideline of 0.1 µg PCBs/g (wet weight) in whole fish, to protect fish-consuming birds and animals, is also based on the Platonow and Karstad (1973) study on minks. As stated above in this document, an application factor of 5 was applied by the IJC to obtain the recommended guideline of 0.1 µg PCBs/g in fish.

To protect human health, Health and Welfare Canada (1975) recommended the maximum residue level of 2 µg PCBs/g (wet weight) in the edible tissue of fish. This guideline by Health and Welfare Canada was adopted in this document.

Swain and Holms (1985) and Nijman and Swain (1990) recommended 0.5 µg/g (wet weight) PCBs as an objective in the tissue of fish caught from the Fraser River and Burrard Inlet. This guideline was adopted from the National Academy of Sciences review (U.S. EPA, 1973), and was based on results with uncertain endpoint. Hence, to protect human health, the guideline of 0.5 µg/g (wet weight) PCBs in fish tissue recommended by Swain and Holms or Nijman and Swain was not considered in this document.

7.4.3 Sediments

Several methods have been proposed in the literature to derive sediment quality guidelines. The sediment criterion recommended in this document is an average value (geometric mean) based on the results obtained using these approaches. The results of equilibrium partitioning (sediment to water and sediment to biota) of PCBs in the environment are shown below. Other approaches include Apparent Effect Threshold (AET), Screening Level Concentration (SLC) and Triad. These approaches employ all relevant data available from the literature. The sediment quality criteria developed by the State of Washington (Table 18) are based on these approaches.

(a) Partitioning of PCBs between Water and Sediment

It was assumed that PCBs sorbed on sediment are inactive and the toxic fraction of PCBs is the one associated with interstitial water. The U.S. EPA (1989) suggested the following relationship between sediment quality criteria (SQC expressed as µg PCB/kg organic carbon) and water quality criteria (WQC expressed as µg PCB /L):

SQC = Kow WQC

where Kow is the octanol-water partition coefficient for PCBs. Given that the average value for Kow for most common PCB formulations (e.g., Aroclors 1016, 1248, 1254 and 1260) is 2.19 x 106 (log Kow = 6.34 - MacKay, 1982) and WQC = 0.0001 µg/L (section 7.3.1), the sediment (for freshwater as well as marine) criterion was calculated to be 0.22 µg PCBs/g organic carbon or 2.2 ng PCBs/g for sediment containing 1% organic carbon.

(b) Partitioning of PCBs between Biota and Sediment

The following results were obtained from partitioning of PCBs between sediment and biota or bio-concentration of PCBs in animal tissues from sediments. The information presented in Tatem (1986) was used in determining the maximum level of PCBs in sediment, using equilibrium partitioning between sediment and benthic organisms. The maximum BAF for freshwater clams (Corbicula fluminea ) exposed to PCBs (Aroclor 1254) in the sediment plus sand mixture (containing 0.111 µg PCB/g dry weight and 0.375% organic carbon) was determined to be 12.5. Assuming the maximum desirable concentration of PCBs in fish/shellfish to be 0.1 µg/g wet weight (see Section 7.4.2), it was calculated that the maximum concentration of PCBs in the sediment should not exceed (0.1 µg/g ÷ 12.5) = 0.008 µg/g dry weight. The results, when expressed in terms of sediment containing 1% organic carbon, would yield the maximum desirable concentration of 0.008 ÷ 0.375 = 0.021 µg PCBs/g-sediment or 21 ng PCBs/g-sediment (containing 1% organic carbon).

(c) Discussion

The two partitioning approaches, (a) and (b) above, yielded sediment PCB criteria of 0.0022 and 0.021 µg PCBs/g-sediment; the geometric mean of the two was determined to be 0.007 µg PCBs/g-sediment. The upper limit (i.e., 0.021 µg PCBs/g-sediment containing 1% organic carbon) of the range obtained using the two partitioning models, is similar to the criteria proposed by CCREM (Chu, 1989), and the objectives proposed for Fraser River and tributaries in British Columbia (Swain and Holms, 1985) (Table 18). (Note that later measurement showed that sediment from the lower Fraser River had an organic carbon content of about 1%.). The National Oceanic and Atmospheric Administration (NOAA) arrived at a guideline (ER-L = 0.05 µg PCB/g sediment - see section 7.1.4), based on all available data, which was about 7 times higher than the geometric mean obtained from the partitioning approach (Long and Morgan, 1990); however, no reference to the organic carbon content of sediment was made in their analysis.

A direct comparison between the sediment criteria obtained above (i.e., in Sections 7.4.3a and 7.4.3b) and those proposed by the State of Washington for Puget Sound (Table 18) is difficult to make since the organic carbon content for sediments was not stated in the Washington State criteria. Note, that the State of Washington used several approaches to derive its criteria (e.g., apparent effects threshold, screening level concentration, and triad).

The criterion of 0.02 µg PCBs/g sediment (freshwater and marine), containing 1% organic carbon, was adopted in this document for two reasons: (a) Sediments containing less than or equal to 0.02 µg PCBs/g (dry weight) are not likely to cause adverse effects on aquatic organisms, and (b) this value represents the mean (geometric) of values obtained using different methods (e.g., partitioning approaches outlined in Sections 7.4.3a and 7.4.3b above, equilibrium partitioning approach used by Chu (Table 18), and the approach used by Long and Morgan (1990) in defining ER-L- see Section 7.1.4).